transport of persistent organic pollutants by ...1 1 transport of persistent organic pollutants by...
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Transport of persistent organic pollutants by microplastics in estuarine 1
conditions 2
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Adil Bakir ab, Steven J. Rowland b and Richard C. Thompson a 4
a Marine Biology and Ecology Research Centre (MBERC) School of Marine Science and 5
Engineering, University of Plymouth, Drake Circus, Plymouth, Devon, PL4 8AA 6
b Biogeochemistry Research Centre, School of Geography, Earth and Environmental sciences, 7
Plymouth University, Drake Circus, Plymouth, Devon, PL4 8AA 8
Corresponding author: [email protected] 9
Room 615, Davy Building, Drake Circus, Plymouth, Devon, PL4 8AA 10
Tel: +44 (0)1752584651 11
Fax: +44 (0)1752232970 12
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Highlights 16
Salinity effect was investigated on sorption/desorption of POPs onto microplastics 17
Little effect on sorption and no effect on their desorption rates was observed 18
Transport of POPs will largely depend on their concentration in each ecosystem 19
A transport model of POPs onto microplastics was proposed 20
Transport followed the order: Phe-PE >> DDT-PVC = DDT-PE >> Phe-PVC 21
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Graphical abstract 24
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Abstract 37
Microplastics represent an increasing source of anthropogenic contamination in aquatic 38
environments, where they may also act as scavengers and transporters of persistent organic 39
pollutants. As estuaries are amongst the most productive aquatic systems, it is important to 40
understand sorption behaviour and transport of persistent organic pollutants (POPs) by 41
microplastics along estuarine gradients. The effects of salinity sorption equilibrium kinetics 42
on the distribution coefficients (Kd) of phenanthrene (Phe) and 4,4’-DDT, onto 43
polyvinylchloride (PVC) and onto polyethylene (PE) were therefore investigated. A salinity 44
gradient representing freshwater, estuarine and marine conditions, with salinities 45
corresponding to 0 (MilliQ water, 690 S/cm), 8.8, 17.5, 26.3 and 35 was used. Salinity had 46
no significant effect on the time required to reach equilibrium onto PVC or PE and neither 47
did it affect desorption rates of contaminants from plastics. Although salinity had no effect on 48
sorption capacity of Phe onto plastics, a slight decrease in sorption capacity was observed for 49
DDT with salinity. Salinity had little effect on sorption behaviour and POP/plastic 50
combination was shown to be a more important factor. Transport of Phe and DDT from 51
riverine to brackish and marine waters by plastic is therefore likely to be much more 52
dependent on the aqueous POP concentration than on salinity. The physical characteristics of 53
the polymer and local environmental conditions (e.g. plastic density, particle residence time 54
in estuaries) will affect the physical transport of contaminated plastics. A transport model of 55
POPs by microplastics under estuarine conditions is proposed. Transport of Phe and DDT by 56
PVC and PE from fresh and brackish water toward fully marine conditions was the most 57
likely net direction for contaminant transport and followed the order: Phe-PE >> DDT-PVC = 58
DDT-PE >> Phe-PVC. 59
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Introduction 61
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Plastics are considered essential in our everyday lives and are used in a wide range of 63
applications, from food packaging, renewable energy and medical devices. World production 64
was 265 million tonnes in 2010, of which 57 million tonnes was produced in Europe (The 65
plastic industry, 2011). The increasing demand for plastics, their low cost of production and 66
high availability mean that end-of-life plastics are accumulating in the environment. The 67
majority of plastic marine litter is believed to come from land-based sources (Sheavly, 2005). 68
The degradation rate of plastic debris in the environment is slow and results in production of 69
small fragments and microplastics. Degradation into small plastic fragments represents an 70
indirect source of microplastics (Barnes et al., 2009; Sivan, 2011) which can also arise from 71
direct sources such as industrial accidental spillages or the release of microbeads used in 72
cosmetics through wastewaters (Browne et al., 2011; Fendall and Sewell, 2009). There is 73
evidence that the abundance microplastics is increasing in the marine environment (Doyle et 74
al., 2011; Thompson et al., 2004) and are potentially bioavailable to a wide range of 75
organisms, for example via ingestion (Browne et al., 2008; Thompson et al., 2009) which 76
have been reported in populations of commercially important fish (Lusher et al., 2013), 77
crustaceans (Murray and Cowie, 2011) as well as seabirds such as the Northern Fulmar (van 78
Franeker, 1985). Laboratory studies have also confirmed that both filter feeding and deposit 79
feeding invertebrates also ingest microplastics (Thompson et al., 2004; Ward and Shumway, 80
2004). Relatively little is known on about the physical (Wright et al., 2013) and toxicological 81
effects (Bakir et al., submitted) of ingestion of microplastics. Microplastics can accumulate in 82
the digestive gland of marine bivalves and microplastics (< 9.6 m) were able to translocate 83
to the haemolymph in the common mussel Mytilus edulis, where they persisted for at least 48 84
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days (Browne et al., 2008). It has also been suggested that ingestion, retention, egestion and 85
possible re-ingestion of microplastics present potential mechanisms for the transport of 86
persistent organic pollutants (POPs), and also for the release of chemical additives from 87
plastics to organisms (Ryan et al., 1988; Tanaka et al., 2013). This will be dependent on the 88
nature of the chemical substance involved, the size of the plastic particle and from the 89
perspective of this study, the surrounding physical and chemical environment. 90
Estuaries are among the most productive marine environments, providing habitats for a wide 91
diversity of seabirds, fish and mammals and are economically important for the exploitation 92
of fish and shellfish (Allen et al., 2006). As microplastics can be carried to the sea via rivers, 93
(Moore et al., 2002) they are likely to be transported to the marine environment via estuaries 94
with the potential to be ingested by estuarine organisms. There is also potential for 95
microplastics to be transported back into estuarine habitats from the sea by tidal flow. Short-96
term transport potential of POPs by microplastics from estuaries to marine waters is under 97
investigation in this study as estuaries represent relatively rapidly changing physical 98
environments in terms of salinity. In contrast, long-term transport potential for POPs by 99
plastic debris, based on mass fluxes, was suggested to be of less importance compared to 100
other pathways such as atmospheric transport and transport of dissolved compounds by ocean 101
currents (Zarfl and Matthies, 2010). 102
Browne et al. (2010) showed that microplastics were abundant in the intertidal zone in the 103
Tamar estuary (UK) and Costa et al. (2011) showed that plastic debris is also found buried in 104
estuarine sediments. Salinity, temperature and the presence of dissolved organic matter 105
appear to be the main parameters governing the solubility of hydrophobic organic compounds 106
such as POPs (Delle Site, 2001). Therefore, it is important to investigate the effects of each of 107
these parameters on the sorption and desorption of POPs by microplastics. The assessment of 108
sorption behaviour of POPs and their transport by microplastics is an important factor in their 109
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environmental risk assessment required to reach Good Environmental Status (GES) as part of 110
the quality descriptor 10 of the Marine Strategy Framework Directive (MSFD 2008/56/EC). 111
MSFD aims to establish “a framework within which Member States shall take the necessary 112
measures to achieve or maintain GES in the marine environment by the year 2020”. 113
The present study investigated the sorption behaviour of phenanthrene (Phe) and DDT onto 114
unplasticised polyvinyl chloride (uPVC) and ultra high molecular weight polyethylene (PE), 115
due to their widespread presence in the marine environment (Browne et al., 2011; Frias et al., 116
2010; Graham and Thompson, 2009; Ng and Obbard, 2006; Thompson et al., 2009), at 117
different salinities to represent transition zones from riverine to marine waters. A transport 118
model, taking into account reported environmental concentrations of the two contaminants 119
for riverine, estuarine and marine waters, is also proposed in order to characterise any 120
potential risks. 121
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Materials and methods 123
Sample preparation and characterisation 124
Unplasticised polyvinyl chloride (uPVC) and ultra high molecular weight (UHMW) 125
polyethylene powder (Goodfellow, Huntington, UK) were sieved to the size range 200 – 250 126
μm to be representative of microplastics found in marine waters (Doyle, 2008; Thompson et 127
al., 2004). PVC and PE were selected as plastics for study due to their widespread presence in 128
the marine environment (Frias et al., 2010; Graham and Thompson, 2009; Ng and Obbard, 129
2006). Seawater was filtered (Whatman membrane filter cellulose nitrate 0.45 m pore size) 130
and autoclaved before use to reduce microbial activity and to remove any suspended 131
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particulate matter (SPM). The salinity range under investigation was 0, 25, 50, 75 and 100 % 132
seawater corresponding to 0 (690 S/cm), 8.8, 17.5, 26.3 and 35 psu (practical scale unit). 133
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Sorption of phenanthrene and DDT to plastics 135
Analysis of samples over a period of 360 hours showed that equilibrium concentrations in 136
seawater and on the plastic were reached after 24 hours (Bakir et al., 2014). Equilibrium 137
sorption time of POPs onto plastics in MilliQ only was also investigated over a period of 360 138
hours. Details of the selected radiolabeled contaminants, suppliers and concentration range 139
are listed in Table 1. The concentration range for Phe (0.6 – 6.1 g L-1) was relevant to 140
environmental concentrations as Phe predominantly enters the marine environment in large 141
pulses through storm waters (Teuten et al., 2007). The concentration range used for DDT in 142
seawater (0.8 – 3.1 g L-1) was lower due to lower concentrations of this legacy pollutant that 143
are typically encountered in the marine environment (Carvalho et al., 2009; Tan et al., 2009). 144
Sorption experiments were conducted in an ISO9001 accredited radioisotope facility at 145
Plymouth University. Either PVC or PE (10 mg) were placed into each of 12 glass centrifuge 146
tubes (50 mL) and an increasing concentration of the POP was added to the walls of the tubes 147
and the solvent allowed to evaporate. 25 mL of seawater were added and the tubes were 148
capped and equilibrated in the dark for 24 hours at 18ºC, with continuous horizontal, rotary 149
agitation (220 rpm). All sorption experiments were conducted in triplicate. 150
The concentration of Phe or DDT was determined in the aqueous and solid phases by 151
counting the β decay from the 14C-Phe and 14C-4,4’-DDT by liquid scintillation counting 152
(LSC) in Ultima Gold (Perkin-Elmer) scintillation cocktail. To measure the aqueous Phe and 153
DDT concentrations, 5 mL of seawater was added to scintillation cocktail and counted by 154
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LSC (Beckman LS 6500 scintillation system). To determine the Phe and DDT concentrations 155
on the sorbents, plastic particles were collected by filtration (Whatman membrane filter 156
cellulose nitrate 0.45 m pore size), added to 5 mL scintillation cocktail and counted directly 157
by LSC. Data indicated that the presence of ≤ 10 mg plastic did not quench the signal or 158
affect the count rate (Teuten et al., 2007). The amount of contaminant in each phase was 159
quantified using a calibration curve prepared by counting known amounts of the contaminant. 160
Total recovery and recovery for each phase are listed in the supplementary information 161
(Table S1). 162
The distribution coefficient, Kd, was calculated using the equation: 163
164
.]/[][ aqesolided CqK Eq. [1] 165
where qe is the amount of contaminant sorbed onto plastic (g kg-1) at equilibrium and Ce is 166
the contaminant concentration in the aqueous phase at equilibrium (g L-1). 167
168
For comparison with the linear model (Kd values), the data were also analysed with the 169
Freundlich model (Eq.2). Freundlich sorption isotherms have been widely used to model 170
binding sorption isotherms for sorbed organic contaminants onto polymers (e.g. Teuten et al., 171
2007). 172
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eFFe CnKq log/1loglog Eq. [2]
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175
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where qe (g kg-1) is the contaminant concentration on the solid phase at equilibrium, Ce (g 176
L-1) is the contaminant concentration in the aqueous phase at equilibrium, KF (L kg-1) is the 177
multilayer adsorption capacity and 1/nF is the Freundlich exponent and an indicator of the site 178
energy distribution of a sorbent (i.e. sorbent heterogeneity increases as n decreases). 179
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Desorption of contaminants from plastic 181
In order to characterise the effect of salinity on the desorption rates of contaminants from 182
plastic, Phe was sorbed to plastics as described above, giving final sorbate concentrations of 2 183
– 4 μg g-1. The filtered solid phase (10 mg) was transferred to a 100 mL amber jar, and 75 mL 184
seawater (full salinity, 35 psu) or 75 mL of MilliQ (0 % salinity, 690 S/cm) were added. 185
The jars were equilibrated in the dark in a water bath with rotary agitation (220 rpm). 186
Aliquots (1 mL) were removed at recorded times (0, 3, 6, 12, 24, 45, 75, 130 and 180 min) 187
and the aqueous concentration of contaminant was determined by LSC. Plastics were allowed 188
to settle for 1 min before sampling. The concentration of contaminant on the solid phase was 189
calculated by difference, taking into account both the amount sorbed to the walls and the total 190
recovery. All kinetic experiments were conducted in triplicate. Pseudo-first order rate 191
analyses were used to determine the rate constant for initial desorption, where the extent of 192
resorption was assumed to be negligible (Teuten et al., 2007). The rate constant (k) for the 193
loss of contaminant from the solid phase was determined from the gradient of plots of 194
ln(Ci/Ci,0) vs. time. 195
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Statistical analysis 197
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Two-way ANOVAs were applied to characterise any significant differences between the 198
distribution coefficients of both DDT and Phe onto PVC and PE. Student-Newman-Keuls 199
(SNK) tests were then used to identify any significant terms. A three factor ANOVA was also 200
applied with salinity, type of plastic and type of pollutant as fixed factors. A multivariate 201
analysis was also performed to compare differences of sorption capacity for different 202
salinities for all POP/plastic combinations. Cochran’s test was used to ensure that the data 203
fulfilled the pre-requisites for parametric analysis and the appropriate data were ln(x+1) 204
transformed. As an additional precaution step for any type I errors, a more conservative p 205
value was selected (p < 0.01) where heterogeneous variances remained after transformation. 206
Analysis was conducted as ANOVA is considered robust to such departure from normality 207
where large data sets are employed (Underwood, 1997). The tests were carried out using 208
IBM® SPSS® Statistics version 20 and are presented in the supplementary information. 209
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Results and discussion 211
We sought to investigate the effects of salinity on the sorption and transport of persistent 212
organic pollutants (POPs) by plastic microparticles in riverine, estuarine and marine systems. 213
As model combinations, sorption equilibrium time of DDT onto PVC and Phe onto PE were 214
investigated in seawater (35 psu) and in MilliQ water over a period of 15 days (Fig.1). 215
Sorption equilibrium was reached in 24 hours and salinity was found to have no significant 216
effect (p > 0.01 and p = 0.229 and p = 0.601) on sorption equilibrium time of Phe onto PE 217
and DDT onto PVC respectively. This indicates that contaminants will reach equilibrium 218
conditions onto plastics relatively quickly in freshwater, brackish or marine waters. 219
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Differences in salinity produced no significant effects on the distribution coefficient values 220
calculated for Phe sorption onto plastics (Fig.2, Table S3). However, significantly higher 221
concentrations of DDT were sorbed onto PVC and PE (p < 0.01, Table S3) in MilliQ water 222
compared at higher salinities. Kd values calculated in MilliQ water for DDT sorption onto 223
plastics were 1.4 and 1.6 times greater for PVC and PE respectively than using seawater (35 224
psu). Kd values calculated for DDT sorption onto PVC than PE showed higher variability than 225
for Phe (Fig.2) and there was a clear trend of decreasing sorption of DDT with increasing 226
salinity. Larger Kd values for DDT sorption onto plastics could be explained by differences in 227
sorption mechanism between DDT and Phe. We previously suggested that DDT sorption onto 228
polymer follows a partition and pore-filling mechanism as opposed to Phe which may only 229
occurs through a pore-filling mechanism, thus explaining higher affinity of DDT for plastic 230
(Bakir et al., 2012; Obst et al., 2011). As aqueous solubility of DDT decreases with an 231
increase in salinity, salinity might be a greater concern for the sorption of DDT onto plastics 232
than for Phe. As salinity had no effect on Phe, following a pore-filling mechanism, salinity 233
may have decreased the partitioning of DDT onto plastics. Solubility of organic compounds 234
decreases with an increase in salinity and salt content in aqueous phase affects partitioning of 235
organic chemicals into other phases (Borrirukwisitsak et al., 2012). 236
The Freundlich model fitted the experimental data well with high correlation coefficients 237
(Fig.S5 and S6 and Table 2). However, differences in salinity had no effects on the sorption 238
capacity of the two plastics as shown by the Log KF values (Fig.S7). This indicates that the 239
sorption behaviour of these two plastics for the two POPs will be influenced by the 240
environmental concentrations of the contaminant, independently of salinity. Log KF values 241
were the lowest for Phe sorption onto PVC, followed by Phe sorption onto PE. Highest 242
values were found for DDT sorption onto PE and PVC. Consequently, PVC and PE have the 243
potential to sorb and transport relatively higher amounts of DDT than Phe. 244
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A study by Karapanagioti et al. (2005), on Phe sorption onto different sorbents, also showed 245
that salinity did not affect the kinetics of sorptive uptake by PE and polyoxymethylene resins 246
(as compared to materials containing organic matter), although the sorption equilibrium point 247
was higher (Karapanagioti et al., 2005). In the present study, salinity did not have any 248
significant effect on the sorption equilibrium kinetics of DDT onto PVC. 249
Sorption behaviour of POPs onto polymers can be compared to sorption onto sediments 250
following a dual-mode model comprising a partitioning and adsorption domain due to the 251
presence of soil organic matter (SOM) (Xing and Pignatello, 1997; Xing et al., 1996; Zhao et 252
al., 2002). However, the effects of salinity appear to be different for marine sediments 253
probably due to the presence of organic carbon which showed a positive correlation between 254
Phe sorption onto sediments and the organic carbon content (Brunk et al., 1996; Yang and 255
Zheng, 2010). Indeed, the sorption behaviour of phenanthrene on marine sediments, showed 256
that the initial rate and equilibrium rate constant of adsorption increased with an increase in 257
salinity (Yang and Zheng, 2010). Another study by Xu et al. (2009) showed that sorption of 258
benzyl butyl phthalate (BBP) on the sediments also increased with salinity. As the sorption 259
potential of a POP is regulated by its solubility, an increase in salinity reduces its solubility 260
and therefore is likely to facilitate its association with sediments (Xu and Li, 2009). This is 261
supported by an earlier study from the same authors on the adsorption behaviour of dibutyl 262
phthalate (DBP) which showed that a decrease in salinity decreased the amount of DBP 263
adsorbed on the sediment (Xu and Li, 2009). Brunk (1997) also suggested that an increase in 264
Ca2+ and Mg2+ might cause the dissolved organic matter (DOM) to precipitate and to coat 265
sediments increasing their organic content and thus their sorption capacity for POPs. The 266
same can therefore be applied for buried plastic materials being coated by organic matter, 267
potentially increasing their sorption capacity for pollutants. 268
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Desorption kinetics of contaminants from plastic 270
Salinity was found to have no significant effects on the desorption rates of sorbed Phe and 271
DDT from PVC and PE (Table 3, p > 0.01, Table S4). Linear plots are shown in the 272
supplementary information (Fig. S8 and S9). Desorption rates for DDT from plastics were 273
much slower than for Phe (p < 0.01, Table S4) but no significant differences were found 274
considering the plastic type (p = 0.087, Table S4). 275
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Transport of persistent organic pollutants (POPs) onto plastics from riverine to marine 277
systems 278
Salinity had little effect, at least for the combinations used herein, on the sorption of 279
microplastics for POPs. If this is generally true for other plastic/contaminant combinations, 280
transport of contaminants from riverine to marine waters is likely to be more dependent on 281
the POP concentration and transport behaviour of microplastics including their residence time 282
in each system (riverine, estuarine or marine) than the type of plastic. Phe and DDT are found 283
in higher concentrations in estuaries compared to riverine and marine waters (Tables 2 and 284
S5). This is not surprising since most anthropogenic organic contaminants (POPs) come from 285
land sources, with estuaries acting as a “sink”, especially for POPs transported by sediments 286
(Ridgway and Shimmield, 2002; Zhou et al., 1999). 287
Using the distribution coefficients (Kd) values measured herein, the amount of Phe and DDT 288
sorbed onto PVC and PE was estimated and plotted for the different systems (rivers, estuaries 289
and marine environment) (Fig.3), using typical aqueous concentrations for these 290
environments (average values shown in Table 4, full concentration listed in Table S5). 291
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Statistical analysis showed significant differences between POP concentrations expected on 292
plastics for the different POP/plastic combinations (p < 0.01). The concentration of POP 293
sorbed onto plastics followed the order: Phe-PE >> DDT-PVC = DDT-PE = Phe-PVC. 294
Significant differences (p < 0.01) were also observed for POP uptake within the different 295
compartments (riverine, estuarine and marine waters) with some exceptions. Statistical 296
analysis showed that sorbed concentration of POPs from plastics was higher for estuaries 297
(salinity 17.5) than for freshwater and marine waters (p < 0.01, Fig.3). 298
Statistical analysis also showed that the concentrations of POPs in each environment were 299
more important (p > 0.01 and p =0.718) than the nature of the plastic (p < 0.01 and p = 0.001, 300
Table S4). Therefore the uptake and transport of POPs by plastics in estuarine conditions will 301
be more related to the input concentrations of the pollutants than the type of polymer present 302
as a carrier. 303
The amount of contaminants sorbed onto plastics is expected to be greater for plastics in 304
estuaries as compared to riverine or marine waters because aqueous POP concentrations are 305
greater in these environments (Fig.3). Sorption equilibrium will be directly dependant on the 306
transport of microplastics from riverine to marine waters, including particle residence time 307
for each system. Our results indicate that sorption equilibrium was reached within 24 hours 308
with no significant effect of salinity. As a result freshwaters, estuaries and marine waters 309
could be considered as single compartments where particles are continuously exposed to 310
organic pollutants at different concentrations. Previous studies have shown that sorption 311
equilibrium for Phe, DDT, perfluorooctanoic acid (PFOA) and di-2-ethylhexyl phthalate 312
(DEHP) onto PVC and PE is reached within 24-48 hours (Bakir et al., 2014). As a result, 313
sorption of POPs onto microplastics is expected to reach equilibrium before their release to 314
the marine environment. 315
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As discussed by Ballent et al. (2012) a series of factors will influence the transport of 316
microplastics from riverine to marine waters. Intrinsic properties of the plastic such as density, 317
shape and size and external factors such as seawater density, seabed topography, flow 318
velocity, turbulence and pressure will dictate the particle retention time in each system (rivers, 319
estuaries and sea). It has also been suggested that transport of microplastics in rivers may be 320
similar to the transport of sediments (Ballent et al., 2012; Browne et al., 2010; Costa et al., 321
2011; Galgani et al., 2000). Fine sediments are carried in suspension (Uncles and Stephens, 322
2010), which is also suspected to be the main transport mechanism for microplastics. Particle 323
transit and retention time will therefore be dependent on several factors such as particle size, 324
and factors specific to the type of estuary such as geographical, meteorological (winds) and 325
morphological settings, stratification and flushing time. Other factors, such as the presence of 326
organic matter or the presence of biofilms, have also been suggested, facilitating the retention 327
of suspended particles in estuaries (Uncles and Stephens, 2010) or by increasing their 328
sorption capacity for pollutants (Brunk et al., 1996). Browne et al. (2010) suggested that 329
small particles of sediment and plastic are both likely to settle slowly from the water-column 330
and are likely to be transported by the flow of water and be deposited in areas where the 331
water movements are slower. 332
Desorption studies indicated that desorption rates will remain consistent independent of 333
salinity. Amounts of desorbed contaminants are expected to be greater in estuaries where 334
particle retention time is expected to be the longest (Ballent et al., 2012). Overall, desorption 335
rates of Phe from plastics in water were found to be faster than for DDT which is suspected to 336
remained sorbed onto plastics. The time required for the concentration of the contaminant 337
onto plastic in estuaries to reach to the concentration of sorbed contaminants in marine waters 338
was estimated using eq.3 and using desorption rates calculated in this study (Fig.S10 and 339
Table S6). 340
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341
)exp(0 ktCC Eq. [3] 342
343
where C0 is the initial POP concentration onto plastic (g g-1) C is the concentration of the 344
POP onto plastic (g g-1) at time t (in days) and k is the desorption rate constant for the first 345
order model (day-1). 346
347
The results indicated that 5 days were required for the concentration of Phe sorbed onto PVC 348
and PE in estuaries to desorb and reach concentration estimated in marine waters (Fig.4). A 349
much longer timescale was required for DDT desorption from plastics with 13 and 15 days 350
for PVC and PE respectively. The potential for microplastics to transport pollutants from 351
estuaries will therefore be dependent on the estuary’s cycle including flushing and particle 352
residence time. Jay et al. (1997) characterised the particle residence time for some estuaries 353
for different locations. Timescales varied from 3 to 9 months for Chesapeake Bay to 0.5 day 354
for Plum Island Sound (Jay et al., 1997). For estuaries presenting a short particle retention 355
time, microplastics are therefore likely to act as vectors for transport of Phe and DDT to 356
marine waters but with limited exposure to organisms. For estuaries with longer retention 357
time, such as Chesapeake Bay, desorption of POPs from plastics could be of concern as well 358
as the exposure of contaminated particles to aquatic organisms. It is however unknown 359
whether the contaminants sorbed onto plastics can then desorb upon ingestion by marine 360
organisms at concentrations which may cause significant harm or ecologically adverse effects. 361
Microplastic ingestion has been documented for a wide range of organisms at many trophic 362
levels (Avery-Gomm et al., 2012; Besseling et al., 2012; Blight and Burger, 1997; Boerger et 363
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al., 2010; Bravo Rebolledo et al.; Browne et al., 2008; Day, 1980; Denuncio et al., 2011; 364
Lusher et al., 2013) with some evidence of trophic transfer of plastics (Farrell and Nelson, 365
2013). Transfer of sorbed contaminants from plastic following ingestion has also been 366
suggested (Tanaka et al., 2013). As estuaries are the most productive ecosystems in the 367
marine environment and due to the relatively long particle retention time in such systems; it 368
is highly likely that a wide range of marine organisms are exposed to contaminated 369
microplastics, but the chemical effects, if any, of this are unknown. Microplastics can enter 370
the marine environment through land-based sources and are exposed to POPs from 371
anthropogenic sources, where sorption takes place. Contaminated particles are then 372
transported into estuaries where they can reside for a relatively long time, depending on the 373
type of estuary. Contaminated particles can also settle, accumulate or being buried onto 374
sediments as well as being deposited on the shoreline (Browne et al., 2011; Browne et al., 375
2010). Man-made operations involving the suspension of sediments, such as dredging, could 376
cause a large release of contaminated particles into the marine environment. 377
Estuaries can also been defined as sink of contaminated particles with their pulse release in 378
marine waters following flushing. Contaminated plastics can also increase in density 379
following biofouling and as a result can be found at the sea surface microlayer (SML), in the 380
water column or deposited onto sediments where they are potentially exposed to a wide range 381
of marine organisms for ingestion. 382
383
Conclusions 384
Increases in salinity had no significant effects on the sorption capacity of PVC and PE for 385
Phe; however sorption capacity for DDT was significantly higher in the absence of salinity 386
suggesting a differential mechanism of sorption for contaminants. The transport capacity of 387
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these plastics for DDT and Phe in estuaries will likely be more related to the aqueous 388
concentration of the contaminants than to the salinity. The transport model proposed showed 389
that the amount of POPs sorbed by plastics in freshwater, to brackish and marine waters 390
followed the order: Phe-PE >> DDT-PVC = DDT-PE >> Phe-PVC. Microplastics have the 391
potential to sorb substantially more contaminants in estuaries due to higher reported 392
concentrations of contaminants than in riverine and marine waters. Sorption equilibrium of 393
POPs onto microplastics is also expected to be reached in each system due to their long 394
particle residence and potential storage in estuarine sediments. As desorption rates of sorbed 395
contaminants were not affected by salinity, amounts desorbed are expected to be correlated to 396
the particle retention time in estuaries and marine waters. The proposed transport model 397
suggested that estuaries can represent an important source for contaminated microplastics 398
with their pulse release to the marine environment through natural (e.g. flushing) and 399
anthropogenic processes (e.g. dredging). 400
This study suggests that the risk assessment for the potential for microplastics to cause harm 401
in the environment, as part of the MSFD, should not be limited to the marine environment but 402
also include estuarine systems which are suspected to represent an important source and sink 403
of contaminated plastics. 404
405
Acknowledgements 406
This study was supported by the Department for Environment, Food and Rural Affairs 407
(DEFRA) in the United Kingdom as part of a larger project investigating the potential for 408
microplastics to cause harm in the marine environment (Grant number ME5416). 409
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19
References 411
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21
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564
565
566
567
22
568
569
Figures 570
571
572
573
Fig.1 Sorption equilibrium time of (a) Phe onto PE and (b) DDT sorption onto PVC in 574
seawater (35 psu) and in MilliQ water (n = 3, ± SD). 575
23
576
577
578
579
580
Fig.2 Distribution coefficients (Kd) calculated from the sorption of Phe onto (a) PVC, Phe 581
onto (b) PE, DDT onto (c) PVC and DDT onto (d) PE for different salinities (n = 3, ± 582
SD). (A, B, C: no significant differences between Kd values; p < 0.01). 583
584
24
585
Fig.3 Potential for microplastics to sorb and transport phenanthrene (Phe) and DDT in 586
riverine, estuarine and marine waters according to reported typical environmental 587
concentrations of Phe and DDT. Calculated amounts sorbed onto plastics via Kd 588
values measured herein. Statistics analysis carried-out between sorbed amounts of 589
contaminants for each POP/plastic combination for the different ecosystems: ** p < 590
0.01; ns: not significant 591
592
593
594
595
596
25
597
Fig. 4 Time required (in days) for contaminants sorbed onto plastics in estuaries to reach 598
concentration of estimated sorbed contaminants onto plastics in marine waters. 599
600
601
602
603
604
605
26
Tables 606
607
Table 1. Pollutants under investigation and associated parameters. 608
609
Polluta
nt
Chemical
structure
Initial
concentrati
on
(g L-1)
Specific
activity
(MBq/mmo
l)
Supplie
r Solvent
MW
(g
mol-1)
Distilled
water
Log
Kow
14C-Phe
0.6-6.1 303.4
Sigma
-
Aldric
h
Methan
ol
178.2
3
600
(22°C)
(EPA,
2013)
4.5
(MacKa
y et al.
(1993))
14C-
DDT
0.8-3.1 370-110
ARC,
Inc.
Ethanol 354.4
9
25
(25°C)
(Registr
y, 2013)
6.36
(Walker
(2008))
610
611
612
613
614
615
616
617
27
Table 2. Freundlich parameters from the sorption of phenanthrene (Phe) and DDT onto PVC 618
and PE for different salinities 619
Salinity range (PSU)
0
(MilliQ)
8.8 17.5 26.3 35
PVC-Phe
Log KF
(L kg-1)
3.3 3.4 3.3 3.3 3.3
nF 0.88 1.01 0.84 0.84 0.87
R2 0.975 0.966 0.954 0.969 0.95
PE-Phe
Log KF
(L kg-1)
4.7 4.8 4.9 5.1 4.9
nF 0.94 0.89 0.86 0.72 0.84
R2 0.995 0.986 0.995 0.979 0.96
PVC-
DDT
Log KF
(L kg-1)
5.8 5.9 5.5 5.3 5.4
nF 0.70 0.63 0.71 0.87 0.82
R2 0.99 0.991 0.988 0.979 0.988
PE-DDT
Log KF
(L kg-1)
5.6 5.6 5.6 5.3 5.9
nF 0.79 0.71 0.72 0.87 0.60
R2 0.973 0.966 0.936 0.993 0.994
620
621
622
28
Table 3. First order rate constants for desorption of phenanthrene (Phe) and DDT from PVC 623
and PE in seawater (full salinity, 39.7) and in MilliQ (n=3, ±SE). 624
Plastic pollutant
Aqueous solution
(salinity, PSU)
Desorption rate
(k, day-1 ± SD)
PVC
Phe
MilliQ (0) 0.73 ± 0.26
Seawater (35) 0.88 ± 0.56
PE
MilliQ (0) 1.15 ± 0.12
Seawater (35) 1.37 ± 0.45
PVC
DDT
MilliQ (0) 0.21 ± 0.01
Seawater (35) 0.26 ± 0.06
PE
MilliQ (0) 0.20 ± 0.04
Seawater (35) 0.23 ± 0.08
625
626
627
628
629
630
631
632
633
634
635
636
637
29
Table 4. Averaged reported environmental concentrations of phenanthrene (Phe) and DDT 638
for riverine, estuarine and marine waters. Full values are shown in Table S5. 639
640
Ecosystems
Contaminant concentration (ng L-1)
Phe DDT
Rivers 243.68 13.97
Estuaries 377.25 120.78
Marine waters 7.5 3.83
641
642
643
644
645
646
647
648
649
650
651
652
653
654
655
656
657
30
Transport of persistent organic pollutants by microplastics in estuarine 658
conditions 659
660
Supplementary information 661
662
Adil Bakir ab, Steven J. Rowland b, Richard C. Thompson a 663
a Marine Biology and Ecology Research Centre (MBERC) School of Marine Science and 664
Engineering, University of Plymouth, Drake Circus, Plymouth, Devon, PL4 8AA 665
b Biogeochemistry Research Centre, School of Geography, Earth and Environmental sciences, 666
Plymouth University, Drake Circus, Plymouth, Devon, PL4 8AA 667
668
Recovery of contaminants 669
Table S1. Total recovery (%) of the contaminants (radioactivity) and recovery in each phase 670
(A: aqueous phase, G: glass walls, S: solid phase, T: total) (averaged values, n=12) 671
672
Salinity (%)
0 25 50 75 100
A G S T A G S T A G S T A G S T A G S T
DDT-PVC 2 3 76 81 2 3 86 91 2 3 72 77 3 2 82 87 2 1 80 83
DDT-PE 2 3 84 89 2 3 86 91 2 3 72 77 3 2 82 87 2 1 83 86
Phe-PVC 51 2 43 96 51 2 45 98 53 2 48 103 54 2 43 99 46 1 40 87
Phe-PE 4 3 77 84 4 6 72 82 46 1 38 85 4 6 73 83 3 8 81 92
673
674
675
31
Sorption equilibrium time 676
Statistical analysis 677
Table S2. Two factor ANOVA and SNK post-hoc test for the effect of salinity on the sorption 678
equilibrium time of (A) Phe onto PE and (B) DDT onto PVC (n=3, ±SD). 679
680
681
682
683
684
685
686
687
688
689
690
691
692
693
694
695
696
697
698
699
700
Source DF MS F P
exposure time (days) 5 0.033 2.85 0.037
salinity 1 0.003 0.28 0.601
exposure time x salinity 5 0.012 1.07 0.401
exposure time (days)
salinity
1 = 2 = 5 = 7 = 10 = 15
MilliQ = seawater
SNK tests
Source DF MS F P
exposure time (days) 5 0.1808 9.61 < 0.01
salinity 1 0.0287 1.52 0.229
exposure time x salinity 5 0.0045 0.24 0.941
exposure time (days)
salinity
SNK tests
MilliQ = seawater
2 < 1 = 5 = 7 = 12 = 15
(A)
(B)
32
Binding sorption isotherms 701
702
Fig.S1 Sorption isotherms of phenanthrene (Phe) sorption onto PVC for different salinities (n 703
= 3, ± SD). 704
705
Fig.S2 Sorption isotherms of phenanthrene (Phe) sorption onto PE for different salinities (n = 706
3, ± SD). 707
33
708
Fig.S3 Sorption isotherms of DDT sorption onto PVC for different salinities (n = 3, ± SD). 709
710
711
Fig.S4 Sorption isotherms of DDT sorption onto PE for different salinities (n = 3, ± SD). 712
713
714
34
Statistical analysis 715
716
Table S3. Multivariate analysis and SNK post-hoc test for the effect of salinity on the 717
distribution coefficient values of POPs onto plastics 718
719
720 721
722
723 724
725
35
Freundlich binding sorption isotherms 726
727
728
729
Fig.S5 Freundlich binding sorption isotherms for phenanthrene (Phe) sorption onto (A) PVC 730
and (B) PE (n = 3, ± SD). 731
732
(A)
(B)
36
733
734
Fig.S6 Freundlich binding sorption isotherms for DDT sorption onto (A) PVC and (B) PE (n 735
= 3, ± SD). 736
737
(A)
(B)
37
738
739
Fig. S7 Log KF values as a function of salinity for Phe and DDT sorption onto PVC and PE 740
(n = 3, ± SD) 741
742
743
744
745
746
747
748
749
750
38
Desorption of contaminants from plastic 751
752
753
754
Fig.S8 Desorption kinetics of phenanthrene (Phe) from (A) PVC and (B) PE in in MilliQ and 755
in seawater (35 psu) (n = 3, ± SD). 756
757
39
758
759
Fig.S9 Desorption kinetics of DDT from (A) PVC and (B) PE in MilliQ and in seawater (35 760
psu) (n = 3, ± SD). 761
762
763
764
765
(B)
40
Statistical analysis 766
Table S4. Three factor ANOVA and SNK post hoc test for the comparison of the effects of 767
salinity on the desorption rates of Phe and DDT from PVC and PE. 768
769
770
771
772
773
774
775
776
777
778
779
Source DF MS F P
POP 1 1.434 66.59 < 0.01
plastic 1 0.071 3.32 0.087
salinity 1 0.015 0.71 0.411
POP x plastic 1 0.097 4.49 0.050
POP x salinity 1 0.003 0.13 0.724
plastic x salinity 1 0.0001 0 0.950
POP x plastic x salinity 1 0.0004 0.02 0.899
POP
Plastic
Salinity
POP x plastic
SNK tests
Phe >* DDT
PE = PVC
MilliQ = seawater
Phe-PVC = Phe-PE >* DDT-PE = DDT-PVC
41
Environmental significance 780
Transport of POPs in estuaries 781
782
Table S5. Reported environmental concentrations of Phe and DDT for riverine, estuarine and 783
marine waters. 784
785
Location
Sampling time
Phe ∑ DDT References
River
St. Lawrence River, Canada 1990 n.a. 1.02 Pham et al., 1993
Henan Reach of the Yellow River, Middle China
2005 - 2006 15.9 – 698.8
n.a. Sun et al., 2009
Gao-Ping River, Taiwan 1999 - 2000 < 20 - 240
n.a. Doong & Lin, 2004
Pearl River Delta, China 2005 - 2006 n.a. 3.89 Guan et al., 2009
Yangtze River 1999
1.37 Sun et al., 2002
Liaohe River 1998
2.77 Zhang and Dong,
2002
Huaihe River
4.45 - 78.9 Yu et al., 2004
Qiantang River 2005
4.9 Zhou et al., 2006
Jiulong River 2000
12.8 Zhang et al., 2001
Red River 1998 - 1999
54.2 Hung and
Thiemann, 2002
Ebro River 1998
3.1 Fernández et al.,
1999
Estuaries
Dalio River Estuary, China 2007 n.a. 1.7 Tan et al., 2009
Hailing Bay, South China 2007 n.a. 0.42 Xing et al., 2009
Minjiang River Estuary 1999 1700 142 Zhang et al., 2003
and 2004
River Thames
1993-1995
< 3 - 304
n.a. Law et al., 1997
River Humber < 8 - 296
River Severn < 3 - 17
River Tamar
River Tees < 8 - 1170
River Tyne < 10 -
90
River Mersey 11 - 73
42
Cadiz Bay n.s. 266-377 n.a. Pérrez-Carrera et
al., 2007
Marine
Gulf of Mexico 2000 n.a. 73-605 Carvalho et al.,
2009
Celtic sea
1993 - 1995
< 6
n.a. Law et al., 1997
Irish sea < 8
Offshore Moray Firth < 8
Offshore Liverpool Bay < 8
Smith’s Knoll < 8
Bering and Chukchi Sea 1999 n.a. 0.013 – 0.123
Zi-wei et al., 2002
Bering sea 1978 n.a. 0.02 Kawano et al.,
1988
South China Sea
1989 -1990 n.a.
0.004 – 0.012
Iwata et al., 1993 Strait of Malacca, Malaysia 0.007
Java Sea, Indonesia 0.006
Phillipines 1999 n.a. 7.4
UNEP, 2002 Thailand 1999 n.a. 3.72
Viet Nam 1999 n.a. 49.27
Singapore 2002 n.a. 0.03 – 2.58 Basheer et al.,
2003
Singapore 2004 n.a. 0.01 – 0.63 Wurl and Obbard,
2005
Antarctic seas 1981-1982 n.a. 0.02 – 0.24 Tanabe et al., 1983
Mediterranean
1977 - 1980
n.a. 1
Subarctic Atlantic n.a. 0.6 Orlova et al., 1983
Tropical Atlantic n.a. 0.7
Northwest Pacific 1976 - 1979 n.a. 0.41 Tanabe and
Tatsukawa, 1980 Indian Ocean 1976 n.a. 0.1
China sea 1977 n.a. 0.08
n.s. not specified, n.a. not assessed 786
787
788
789
790
791
792
793
43
Desorption of sorbed contaminants 794
795
796
797
Fig.S10 Estimated time required for concentration of POPs onto plastics sorbed in estuaries to reach 798
the concentration of sorbed POPs in marine waters. 799
800
Table S6. Time required (in days) for estimated sorbed concentration of POPs onto plastics in 801
estuaries to reach estimated concentration in marine waters 802
803
POP/plastic Time (days)
Phe-PVC 5
Phe-PE 5
DDT-PVC 13
DDT-PE 15
804
805
44
806
807